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Subject: Synthesis of iron oxyhydroxide-coated rice straw (IOC-RS) and its application in arsenic(V) removal from water

Because of the recognition that arsenic (As) at low concentrations in drinking water causes severe health effects, the technologies of As removal have become increasingly important. In this study, a simplified and effective method was used to immobilize iron oxyhydroxide onto a pretreated naturally occurring rice straw (RS). The modified RS adsorbent was characterized, using scanning electron microscope, Fourier transform infrared spectroscopy, thermogravimetric analyzer, and surface area analyzer. Experimental batch data of As(V) adsorption were modeled by the isotherms and kinetics models. Although all isotherms, the Langmuir model fitted the equilibrium data better than Freundlich and Dubinin–Radushkevich models and confirmed the surface homogeneity of adsorbent. The iron oxyhydroxide-coated rice straw (IOC-RS) was found to be effective for the removal of As(V) with 98.5% sorption efficiency at a concentration of <50 mg/L of As(V) solution, and thus maximum uptake capacity is ∼22 and 20 mg As(V)/g of IOC-RS at pH 4 and 6, respectively. The present study might provide new avenues to achieve the As concentrations required for drinking water recommended by the World Health Organization.

INTRODUCTION

Arsenic (As) is a toxic metalloid which can pollute water, land, crops, and the environment at large, ultimately affecting human health (Zhao et al. 2010a, b). A high level of As (>10 mg/L) may cause skin lesions, rhagades, damage mucous membranes, digestive, respiratory, circulatory, nervous system, and moreover it is associated with skin, liver, and lung cancers (Choong et al. 2007; Nguyen et al. 2009; Lizama et al. 2011; Bulut et al. 2014). During the last 20 years, naturally occurring As has been found to be widespread in natural water in many countries around the world, especially in Bangladesh, India (West Bengal), China (including Inner Mongolia), Vietnam, Argentina, and some parts of West Africa due to their exposure to high As drinking water sources (Zhang et al. 2003; Smedley et al. 2007; Pehlivan et al. 2013). Works on ground waters from different parts of Burkina Faso have shown that the Yatenga province population has been exposed to As contaminated water with over 0.5–1,600 mg/L (Smedley et al. 2007; Barro-Traoré et al. 2008).

Arsenic cannot be destroyed in the environment; however, it can be changed to different forms and accumulated in different biota and environmental media. Various technologies available for removal of As from contaminated water are based mainly on six principles: (i) oxidation and filtration; (ii) biological oxidation: oxidation of As(III) to As(V) by micro-organisms and then removal of As(V) by iron and manganese oxides; (iii) co-precipitation: oxidation of As(III) to As(V) by adding suitable oxidizing agent followed by coagulation, sedimentation, and filtration; (iv) adsorption: activated alumina, activated carbon, iron-based sorbents, zero-valent iron, and hydrated iron oxide, etc.; (v) ion-exchange through suitable cation and anion exchange resins; and (vi) membrane technology: reverse osmosis, nanofiltration, and electrodialysis (Mondal et al. 2006; Lizama et al. 2011; Jain & Singh 2012). Among these, a few treatment technologies are efficient but expensive whereas some are cheaper but not efficient. Because of the simplicity of the adsorption process, it has become one of the most promising and applied methods in As removal (Kanel et al. 2006; Bulut et al. 2014). The most commonly employed adsorbents used for removal of As include magnetic iron oxide nanoparticles (Song et al. 2013), activated alumina (Singh & Pant 2004), resin-MnO2 (Lenoble et al. 2004), TiO2 (Dutta et al. 2004), nanostructured ZrO2 spheres (Cui et al. 2013), ores (Chakravarty et al. 2002; Zhang et al. 2004), zeolites (Elizalde-Gonzalez et al. 2001), activated carbon (Kalderis et al. 2008), functionalized graphene (Mishra & Ramaprabhu 2011), p(4-vinylpyridine)-based hydrogels (Sahiner et al. 2011), iron-modified: resin (Rau et al. 2003), sponge (Munoz et al. 2002), and sand (Thirunavukkarasu et al. 2003). Unfortunately, most of these materials are considered as expensive adsorbent in many countries.

Rice straw (RS) is one of the most abundant natural sources in the world. Its annual production is about 731 million tonnes, which is distributed in Asia (667.6 million tonnes), America (37.2 million tonnes), Africa (20.9 million tonnes), Europe (3.9 million tonnes), and Oceania (1.7 million tonnes) (Kim & Dale 2004; Binod et al. 2010; Hsu et al. 2010). It is a kind of lignocellulosic biomass which contains about 32–47% cellulose, 19–27% hemicellulose, and 5–24% lignin (Karimi et al. 2006; Wattanasiriwech et al. 2010). This material is one of the most agriculture wastes listed in Burkina Faso. In this context, we decided to probe the adsorption capabilities of RS-based matrixes toward As(V).

This paper presents the synthesis of an inorganic–organic hybrid adsorbent of iron oxyhydroxide-coated rice straw (IOC-RS) by RS: (i) pretreatment in sulfuric acid and sodium hydroxide solutions; and (ii) impregnation in ferric nitrate and sodium hydroxide solutions. This work focused on investigating how various experimental parameters influence As(V) adsorption. These parameters included: pH, As(V) initial concentration and the interference of dissolved NaNO3, MgSO4, and Na3PO4.

EXPERIMENTAL

Chemicals

H2SO4 and NaOH used for RS pretreatment were purchased from Merck. Fe(NO3)3 procured from Merck was used for biosorbent modification. A Titrisol ampule with As2O5 in H2O used for batch sorption experiments was purchased from Merck, Germany. As(V) stock solution of 1,000 mg/L As(V) was prepared by transferring the Titrisol ampule content with As2O5 in H2O (Merck, Darmstadt, Germany) into a 1 L volumetric flask, which was filled up to the mark at 20 °C and stored at 16 °C. Diluted solutions of As(V) were prepared daily before starting the batch studies. NH4OH and HCl purchased from Merck, Germany, were used to adjust the pH solutions. MgSO4·7H2O (Fluka, Seelze, Germany), Na3PO4·12H2O (Sigma-Aldrich, Seelze, Germany) and NaNO3 (Merck) were used for studying the effects of ionic strength. KI used for reduction of As(V) to As(III), was obtained from Roth, Germany. NaBH4 (Fluka, Seelze, Germany), NaOH (Merck, Darmstadt, Germany), and HCl (Merck, Darmstadt, Germany) were used for hydride generation (HG). All reagents used throughout this work were of analytical grade. All glassware was cleaned with diluted nitric acid solution and rinsed with deionized water. Pure water was obtained with a water system filtration (Seralpur, Pro90C).

Pretreatment of raw RS

RS was collected from local farm near the Ouagadougou area, Burkina Faso. The straws were cut into small pieces with lengths of ≤6 cm and then washed three times with tap water. The air-dried material (at 60 °C) was made into powder in an electrical ball mill (BLB Braunschweig) and sieved in an electrical sieving machine (Retsh, West of Germany). The particles of sizes between 0.125 and 0.200 mm were rewashed thoroughly with deionized water to remove the fine particles, and dried in a hot-air oven (at 70 °C) for 24 h. The air-dried and powdered RS (60 g) was successively treated with: (i) 2 mol/L H2SO4 (1/1 w/w of dry matter, at 80 °C for 30 min) to remove starch, proteins, and sugars; and (ii) 0.5 mol/L NaOH (ratio straw/sodium hydroxide = 5, stirring for 24 h at 22 °C) to remove the low molecular weight lignin compounds after filtration. The substrate was thoroughly washed, and then the material was air-dried in an oven at 60 °C for 24 h. The dried material was stored in a vacuum desiccator.

Preparation of the sorbent

Twenty-five grams of the dry acid and alkali-treated RS was soaking with 100 mL deionized water for 2 h. The soaked substrate was crushed in a porcelain mortar to get a solid paste. Coating was performed by reacting the solid pasty material with 800 mL of ferric nitrate Fe(NO3)3 solution (0.05 mol/L) in a 2 L flask. Sodium hydroxide (1 mol/L) was slowly added into the mixture, under continuous stirring, until the pH rose to a value between 2.8 and 3.2. The loading process was left for 24 h, and the pH was checked and readjusted with acid (0.1 mol/L HCl) and base (0.1 mol/L NaOH) solutions during the process. The sorbent was then filtered, washed with deionized water several times, dried in an oven at 50 °C for 24 h and stored in desiccators.

Methods of characterization

The thermogravimetric analysis (TGA) of RS and IOC-RS were analyzed using a SETSYS Evolution TGA 16/18 Setaram Instrumentation. The ramping rate was 10 °C/min up to 900 °C in a nitrogen environment. The scanning electron microscope (SEM) was performed on JEOL JSM-6480 microscope to collect the SEM images of the adsorbents. The Fourier transform infrared spectroscopy (FTIR) was carried out on a Bruker Tensor 27 spectrometer, Diamant attenuated total reflectance. The spectra were recorded in the region of 4,000–520 cm−1 for 32 scans. The specific surface area was determined by N2 adsorption at −196 °C using an ASAP 2020 Micromeritics instrument and the Brunauer–Emmett–Teller method (Sing et al. 1985). The zeta potential measurement of the sorbent was performed by mixing 0.2 g of sample with 50 mL of deionized water at 22 ± 2 °C. The pHi (initial pH) values of the solution were adjusted roughly from pH 4 to 10 by a pH meter, via addition of 0.1 M HCl or NH4OH. The mixture was shacked to the equilibrium time of 4 h. After completion of the equilibration time, the admixture was filtered and the pHf (final pH) values were measured. The difference between pHi and pHf values (ΔpH = pHf − pHi) was plotted versus the pHi. The pH value at the point of zero charge (pHpzc) of the sorbent was determined from the point of intersection of the resulting curve, at which ΔpH = 0. For As(V) concentrations access, HG attached with atomic absorption spectrometry with a Zeeman correction (AAS-Hitachi Z-2000) was used, with a calibration range from 1 to 20 μg/L. Arsine obtained from As(III) gives the best signal; therefore, all As(V) in the sample was previously reduced to As(III). The reduction was carried out with the optimized protocol: 30% HCl (2.5 mL) and 20% KI (2.5 mL) solution as a reduction agent were added to all sample solutions (10 mL), and a waiting time of 1 h allowed for complete reduction of As(V); then the whole were made up to 25 mL before analysis. For HG, the following solutions were used: (i) 1.2 M HCl; and (ii) NaBH4–NaOH solution (solute 10 g NaBH4 in 1 L of H2O Seralpure by adding 4 g of NaOH); the solution was prepared immediately before use. The principle is to volatilize As from As(III) to arsine AsH3(g), due to a reaction of nascent hydrogen on As. This nascent hydrogen comes from the decomposition of BH4− ions.

Batch sorption experiments

Batch experiments were performed in plastic bottles (50 mL) by adding the sorbent in 50 mL of aqueous As(V) solution of the desired initial pH. For all experiments, the initial pH of the As(V) solution was controlled every 30 min with a digital pH meter by adding 0.1 M HCl and/or NH4OH solution as required. The bottles were gently agitated with a rotary shaker (Retsch, Berlin, Germany) at 120 rpm. The sorbent was separated by filtration with cellulosic acetate film (pore size 0.2 μm) and the remaining As was analyzed using HG atomic absorption spectrometry. Each experiment was replicated three times at the desired initial conditions and the mean (average) values are taken. The amount of As(V) adsorption at equilibrium, qe (mg/g), was calculated using the following equation:

1

where C0 and Ce are the As concentrations (mg/L) initially and at equilibrium, respectively. V is the volume of the As(V) solutions (L), and W is the weight of sorbent (g). qe (mg/g) is the adsorption capacity at equilibrium.

Effect of solution initial pH on As(V) sorption

The effect of solution pH was carried out by adding 0.2 g of sorbent in 50 mL of As(V) solution of 50 mg/L initial concentration at different pH values (2.0–10.0). The mixture was gently agitated with a rotary shaker (Retsch, Berlin, Germany) for 8 h at a temperature of 22 ± 2 °C.

Effect of initial concentration and ionic strength on As(V) sorption

The influence of initial solution concentration on As(V) sorption with sorbent was carried out by adding 0.2 g of sorbent into 50 mL of As(V) solution of 10–300 mg/L initial concentration. The pH values were adjusted at 4 and 6, and the mixtures were shaken for 24 h at 22 ± 2 °C. The As sorption experiments were repeated with solution containing the mixture ions of dissolved NaNO3, MgSO4, and Na3PO4, usually present in water with As(V) solution of 50 mg/L.

RESULTS AND DISCUSSION

Sorbent characterization

The SEM pictures (Figure 1) reveal the surface textures and porosities of RS and IOC-RS. It was found that after the surface of RS had been coated by the iron oxyhydroxide in the iron reclaim system, the apparent color of RS changed from dark yellow to dark orange (red), and the surface morphology became much finer and smoother. On the other hand, the surface area of the IOC-RS (7.86 m2/g) was also different from that of RS (2.85 m2/g).

RS is a lignocellulosic (cellulose, hemicellulose, and lignin) and silica matrix, which likely consists of alkene, esters, aromatic, ketones, and alcohols with different oxygen-containing functional groups. The FTIR spectrum of RS (Figure 2(a)) showed bands at 897 cm–1, 1,060 cm–1, and 1,380 cm–1, which are ascribed to β-glycosidic linkages, C–O–C stretching and O–H bending, respectively (Sun et al. 2004). The bands observed at 3,332 cm–1 and 1,735 cm−1 can be assigned to O–H stretching and C=O stretching, respectively. The presence of C=C groups is reflected by vibration bands at 1,606 and 1,515 cm−1 due to aromatic rings of lignins (Fernandez-Bolanos et al. 1999). On the other hand, the adsorption peaks at 1,645 cm−1 and 796 cm−1 are, respectively, assigned to the H–O–H bending mode (Liu et al. 2006) and the symmetric Si–O–Si stretching vibration (Farook et al. 2006). The loaded RS spectrum (Figure 2(b)) shows some changes mainly in the decrease of the peak intensities at 1,735 and 1,645 cm−1, which can be explained by the complexation of Fe3+ ions with carboxylate and hydroxyl groups in the matrix (Pehlivan et al. 2013). The decrease of peaks observed at 1,606, 1,515, and 796 cm−1 is due to a partial removal of lignins and silica during the pretreatment with NaOH (Sun et al. 2004; Liu et al. 2006). The strong band between 600 and 520 cm−1 belongs to the stretching mode Fe–O (Correa et al. 2010; Bordoloi et al. 2013; Bulut et al. 2014).

The thermal degradation (Figure 3) of RS and IOC-RS showed that there was a similar weight loss between 150 and 550 °C due to degradation of lignocellulose polymers. The minor weight loss of ∼1.5% was first observed around 160 °C in the RS, which may be due to the decomposition of volatile matters such as low molecular weight sugars. Following this, a substantial weight loss by ∼50% took place from 200 to 400 °C. This was expected to be the decomposition of hemicellulose and cellulose (Wattanasiriwech et al. 2010); the overlapping peaks in the differential thermogram (DTG) are, respectively, 322 and 366 °C. Another important loss of ∼14% due to the decomposition of lignin occurred between 350 and 540 °C. For IOC-RS, the result clearly shows that the thermal stability changed by modification of RS with iron. The DTG shows one peak at 354 °C, suggesting that it was likely that hemicellulose and cellulose simultaneously decompose within this temperature. The percentage weight loss of the residue after heating up to 700 °C was ∼4% higher for IOC-RS when compared to unmodified RS. This implies that there was presence of iron in IOC-RS which was not degraded when the samples were heated up to 700 °C. In addition, it was found that ash content in the pretreated RS was only 3.53% whereas that in the IOC-RS was increased to 22.84%; this higher percentage of residual ash in IOC-RS is assumed to be from iron(III)–lignocellulose complex. The form of iron in the IOC-RS was iron oxyhydroxide (–FeOOH), as previously reported (Pehlivan et al. 2013).

The pHzpc (zero proton charge (zpc)) of an adsorbent is a very important characteristic that determines the pH at which the sorbent surface has net electrical neutrality, and at which value the acidic or basic functional groups no longer contribute to the pH of the solution (Wan Ngah & Hanafiah 2008). Figure 4 shows a plot of the zeta potential (ΔpH) of sorbent versus initial pH (pHi). The values of the zeta potential in all suspensions decreased as the pH was increased. According to the zeta potential curve of sorbent, it was found that the pHzpc of IOC-RS was about 5.9. This result is in agreement with the pHzpc of Fe(III)-coated rice husk (Pehlivan et al. 2013). Based on the pHzpc value, it can be deduced that the IOC-RS surface charge is globally positive, as the solution pH is less than 6. On the contrary, with a solution pH higher than the pHzpc, the surface is negatively charged and there will be an electrostatic repulsion between an anionic species and the surface of IOC-RS.

The effect of solution initial pH on As(V) sorption

To determine the optimum pH for adsorption of As over RS and IOC-RS, the uptake of As(V) was studied at pH range of 2–10 and the removal data are shown in Figure 5. Results from the present study clearly show that As(V) adsorption by RS is not effective whatever the solution pH; while the percent removal of As(V) by IOC-RS was reduced from 99.6 to 64.3% as pH shifted from 2 to 10. Optimal As(V) adsorption by IOC-RS was found in the range of pH 2–4 in which the removal rate was above 95%. The pH affects significantly the speciation of As(V) in solution and the surface charge of the solid particles. As(V) species and their corresponding stability pH values are H3AsO4 (pH <2), H2AsO4− (pH 2–7), HAsO42− (pH 7–11), and AsO43− (pH >12). As pH increased from 4 to 10, the amount of multivalent species (HAsO42− (pH 5–9) and AsO43− (pH >9)) increased and the two species were not preferably adsorbed by IOC-RS in comparison with H2AsO4−.

It was previously reported that the main factors governing the adsorption of As species are (i) the electrostatic interaction between iron oxyhydroxide sites of the adsorbent surface and (ii) the anionic As(V) species (Pehlivan et al. 2013). As the equilibrium pH increased from lower pH to pHzpc, the decreased percentage removal of As(V) was attributed to the decreasing electrostatic attraction between the surface of iron oxyhydroxide loaded in IOC-RS and anionic multivalent HAsO42− and AsO43− species. Over the pHzpc value, the surface sites of IOC-RS were negatively charged and inappropriate for adsorbing anionic arsenate species. The decrease in arsenate adsorption at pH > pHzpc may have been due to the increase in the negative charge density at the surface of the adsorbent, and to the increase in the number of OH− ions in the solution, in competition with the anionic arsenate species for adsorption.

The effect of initial concentration and isotherm study

To be able to estimate maximum capacities of adsorbents, it is necessary to know the quantity of adsorbed As(V) as a function of As(V) concentration in solution. The As(V) adsorption isotherms for the sorbent are shown in Figure 6. The sorption data obtained were analyzed by fitting the Freundlich and Langmuir isotherm models (Ranjan et al. 2009; Mishra & Ramaprabhu 2011; Maji et al. 2013). Table 1 gives a full overview of the Freundlich and Langmuir adsorption isotherm parameters.

The Freundlich isotherm is most frequently used to describe the adsorption of heavy metal ions in solution. The Freundlich isotherm assumes that the uptakes of metal ions occur on a heterogeneous surface by multilayer adsorption. The equilibrium data were analyzed using the following Freundlich equation:

2

where 1/n is the intensity of adsorption, Kf is the adsorption capacity, qe is the amount of As(V) adsorbed per unit amount of the adsorbent (mg/L), and Ce is the equilibrium concentration in solution. The Kf and 1/n values are calculated from the linear plot of ln qe versus ln Ce (Table 1). The 1/n value was between 0 and 1 indicating that the sorption of As(V) using understudy sorbent material was favorable at the studied conditions. However, the R2 value was found to be >0.9784, indicating that the Freundlich model was applicable for the relationship between the amounts of As(V) ions sorbed and its equilibrium concentration in the solution.

The Langmuir model has eventually been empirically the best known of all sorption isotherms used since it contains the two useful and easily imaginable parameters (b and Q) which are more easily understandable. A basic assumption of this theory is that sorption takes place at specific homogeneous sites within the sorbent. This model can be written in linear form

3

where Ce is the equilibrium concentration of As(V) (mg/L) in solution, qe is the amount of As(V) (mg/g) adsorbed per unit mass of adsorbent, and Q is the monolayer sorption saturation capacity (mg/g). b (L/mg) is a constant related to the affinity of binding sites or bonding energy.

A plot of Ce/qe versus Ce is given in a straight line with its slope of 1/Q and intercept of 1/Qb and the results are shown in Table 1. According to the coefficients of correlation obtained (R2 > 0.9961), the sorption of As(V) ions onto sorbent material, are fitted well to the Langmuir model. The maximum adsorption capacity of sorbent material for As(V) removal at pH 6 and 4 were found to be 19.96 mg/g and 21.739 mg/g, respectively (Table 1). Recently, several studies related to As(V) ions adsorption from water used for drinking and irrigation of crops have been carried out. Table 2 compares As(V) adsorption capacities of iron-based material adsorbents reported in previous studies. As seen, iron(III)-loaded resin has a high adsorption capacity for As(V) at pH < 2. In comparison to these approaches, the IOC-RS material (at pH 6) reported herein achieved better results than the amorphous iron oxide and ferric oxyhydroxides-modified sawdust, displaying at least five-fold better removal efficiency than the iron oxide nanoparticles and iron(III)-coated rice husk and a higher efficiency compared to the iron ores, iron oxide-coated sand (IOCS) and zeolite (ICZ). On the other hand, it was also found that the adsorption capacities of IOC-RS carried out at pH 4 and 6 are, respectively, comparable to those of iron-coated chitosan flakes (ICF) (pH 7) and iron(III)-loaded sponge (pH 9). The content of iron in the matrix is a crucial factor which impacts on the As adsorption capacity. In our case, there was a significant increase of ash content, which indicates high amounts of iron adsorbed on the surface of the RS substrate compared to the reported rice husk. This could relate to the chemical and physical properties of the feedstock. Rice husk is made of hard materials, including opaline silica and lignin, and has a more recalcitrant cell wall structure than RS, which contains a lesser amount of lignin (Boonmee 2012). In a future study, the correlation between the amount of iron-coated RS and the As adsorption capacity will be investigated. Moreover, it was found that IOC-RS could be regenerated by HCl and NaOH treatment. The highest recovery of >88% was achieved with 1 M NaOH.

To determine that the nature of sorption processes is physical or chemical, the equilibrium data were also modeled using the Dubinin–Radushkevich (D–R) isotherm equation (Namasivayam & Sureshkumar 2008; Baig et al. 2010) as follows:

4

where

5

where qe is the amount of As ions adsorbed on per unit weight of sorbent material (mol/g), and Ce is the concentrations at equilibrium (mol/L). Xm is the maximum sorption capacity (mol/g), β is the activity coefficient (mol2/kJ2) related to sorption mean free energy (kJ/mol), and ε is the Polanyi potential, where R (8.32 J/mol/K) is the gas constant and T (K) is the absolute temperature. The constant β and Xm (Table 1) were estimated from slope and intercept of the plot of ln qe against ɛ2. The R2 value shows that the present data describe well the D–R equation.

The mean free energy of adsorption (E), defined as free energy change when 1 mol of As(V) ion is transferred from infinity in solution to the surface of IOC-RS, can be calculated using the value of β, according to the following equation:

6

It was reported that when the magnitude of E is: (i) less than 8 kJ/mol, physical adsorption is the major process; (ii) between 8 and 16 kJ/mol, the adsorption behavior is dominated by ion-exchange; and (iii) in the range of 20–40 kJ/mol, chemisorption is predominant in the adsorption procedure (Han et al. 2013). The estimated values of E (Table 1) obtained at pH 4 and 6 were respectively 22.36 and 25 kJ/mol, suggesting that the sorption of As(V) ion on the surface of IOC-RS took place with chemisorption mechanism.

Effect of concomitant ions

The sorption of As(V) in the presence of common anions ions may be affected due to precipitation, complex formation or competition for sorption sites. Interference of anions on the sorption of As(V) onto sorbent was carried out with As(V) solution containing 50 mg/L of anion (i.e., NO3−, SO42−, or PO43−) ions, at pH 4. It was observed that except for PO43−, other ions, e.g., NO3− and SO42− have no significant interference with sorption of As(V) ions. The decrease of percentage removal of As(V) in the presence of PO43− was 6%.

CONCLUSION

In this work, an inorganic–organic hybrid IOC-RS was used as adsorbent for the removal of As(V) from aqueous solution. The major conclusions based on the experimental study were as follows:

The pH, contact time, initial concentration, and dissolved Na3PO4 on the adsorption, significantly affect the As(V) adsorption capacity of IOC-RS.

The isotherm modeling revealed that the Langmuir equation could better describe the adsorption of As(V) on the IOC-RS as compared to Freundlich and D‒R models. The maximum adsorption capability of IOC-RS reached ∼22 mg As(V)/g of IOC-RS at pH 4.0 and 22 °C. In addition, the adsorption energy indicates that the adsorption process is dominated by the chemisorption.

The removal capacity Q of IOC-RS sorbent for As(V) ions was found to be higher than that of the majority of other iron oxide sorbents reported in the literature. Therefore, it can be stated that this sorbent has significant potential for the removal of As(V) ions from natural water.

ACKNOWLEDGEMENTS

This investigation was performed at the Guest Chair within the project ‘Exceed – Excellence Center for Development Cooperation – Sustainable Water Management in Developing Countries’ at the Technische Universitaet Braunschweig, Prof. Pehlivan being the visiting professor, and Ms Tran and Dr Ouédraogo being the international exchange staff members. The Exceed Project is granted by the German Federal Ministry for Economic Cooperation and Development (BMZ) and German Academic Exchange Service (DAAD), and their financial support we gratefully acknowledge.